v>EPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/600/R -92/081
September 1993
Methods for Aquatic
Toxicity Identification
Evaluations
Phase III Toxicity
Confirmation Procedures for
Samples Exhibiting Acute and
Chronic Toxicity
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EPA/600/R-92/081
September 1993
Methods for Aquatic
Toxicity Identification Evaluations
Phase III Toxicity Confirmation
Procedures for Samples Exhibiting
Acute and Chronic Toxicity
by
D. I. Mount1
T. J. Norberg-King2
With Contributions from:
G. T. Ankley2
L. P. Burkhard2
E. J. Durhan2
M. K. Schubauer-Berigan1
M. T. Lukasewycz1
'AScI Corporation - Contract No. 68-CO-0058
*U.S. Environmental Protection Agency
Previous Phase III Methods
by D.I. Mount
EPA-600/3-88/036
National Effluent Toxicii Assessment Center
Technical Report 02-93
Environmental Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Duluth, MN 55804
Printed on Recycled Paper
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Disclaimer
This document has been reviewed in accordance with U.S. Environmental
Protection Agency Policy and approved for publication. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use.
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Foreword
This Phase III document is the last in a series of guidance documents
intended to aid dischargers and their consultants in conducting aquatic organism
toxicity identification evaluations (TIEs). TIEs might be required by state or federal
agencies as the result of an enforcement action or as a condition of a National
Pollutant Discharge Elimination System (NPDES) permit. These documents should
aid individuals in overseeing and determining the adequacy of effluent TIEs as a part
of toxicity reduction evaluations (TREs).
There are two major reasons to require the confirmation procedures. First the
effluent manipulations used in Phase I characterizations (EPA, 1988; EPA, 1991 A;
EPA, 1992) and Phase II identifications (EPA, 1989A; EPA, 1993A) might (with some
effluents) create artifacts that might lead to erroneous conclusions about the cause
of toxicity. Therefore in Phase III confirmation steps, manipulations of the effluent are
avoided and/or are minimized, therefore artifacts are far less likely to occur. Some-
times, toxicants will be suspected through other approaches (such as the treatability
route) which on their own are not definitive and in these instances, confirmation is
necessary. Secondly, there is the probability that the substances causing toxicity
might change from sample to sample, from season to season or some other
periodicity. As toxicity is a generic measurement, measuring toxicity cannot reveal
variability of the suspect toxicant whereas the Phase III confirmation procedures are
designed to indicate the presence of variable toxicants. Obviously, this crucial
information is essential so that remedial action may be taken to remove toxicity.
Confirmation, whether using the procedures described in this document or
others, should always be completed because the risk is too great to avoid or eliminate
this step. Especially for discharges where there is little control over the influent or for
discharge operations that are very large or complex, the probability that different
constituents will cause toxicity over time is great. Most of the approaches in Phase III
are applicable to chronically toxic effluents and acutely toxic effluents.
In this confirmation document, guidance is included when the treatability
approach (EPA, 19898; EPA, 1989C) is taken. Use of the treatability approach
requires confirmation as much as or more than the toxicant identification approach
(Phase II). The reader is encouraged to use both the acute Phase I characterization
(EPA, 1991 A) and the chronic Phase I characterization (EPA, 1992) documents for
details of quality assurance/quality control (QA/QC), health and safety, facilities and
equipment, dilution water, sampling and testing. The TIE methods are written as
general guidance rather than rigid protocols for conducting TIEs and these methods
should be applicable to other aqueous samples, such as ambient waters, sediment
elutriate or pore waters, and leachate.s
lii
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Abstract
In 1989, the guidance document for acutely toxic effluents entitled Methods
for Aquatic Toxicity Identification Evaluations: Phase III Toxicity Confirmation Proce-
dures was published (EPA, 19890). This new Phase III manual and its companion
documents (EPA, 1991 A; EPA, 1992; EPA, 1993A) are intended to provide guidance
to aid dischargers in confirming the cause of toxicity in industrial and municipal
effluents. The toxicity identification evaluation (TIE) starts with a characterization of
the effluent toxicity using aquatic organisms to track toxicity; this step is followed by
identifying a suspect toxicant(s) and then confirming the suspect toxicant as the cause
of toxicity.
This Phase III confirmation document provides greater detail and more
insight into the procedures described in the acute Phase III confirmation document
(EPA, 1989D). Procedures to confirm that all toxicants have been correctly identified
are given and specific changes for methods applicable to chronic toxicity are included.
Adifficult aspect of confirmation occurs when toxicants are not additive, and therefore
the effects of effluent matrix affecting the toxicants are discussed. The same basic
techniques (correlation, symptoms, relative species sensitivity, spiking, and mass
balance) are still used to confirm toxicants and case examples are provided to
illustrate some of the Phase III procedures. Procedures that describe the techniques
to characterize the acute or chronic toxicity (EPA, 1988) and to identify (EPA, 1989A)
toxicants have also been rewritten (EPA, 1991 A; EPA, 1992; EPA, 1993A).
IV
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Contents
Page
Foreword iii
Abstract iv
Tables vi
Figures vi
Acknowledgments vii
1. Introduction 1-1
2. Correlation Approach 2-1
2.1 Correlation 2-1
2.2 Correlation Problems Caused by Matrix Effects 2-4
3. Symptom Approach 3-1
4. Species Sensitivity Approach 4-1
5. Spiking Approach 5-1
6. Mass Balance Approach 6-I
7. Deletion Approach 7-1
8. Additional Approaches 8-I
9. Hidden Toxicants 9-1
10. Conclusions 1 o-1
11. When the Treatability Approach Has Been Used 11-1
12. References 12-1
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Tables
Number Page
6-1. Comparison of effluent toxicity and toxicity measured in effluent
fraction add-back tests
Figures
Number Page
2-1. Correlation of toxic units (IDs) for an effluent and one suspect toxicant
in a POTW effluent 2-2
2-2. Correlation of toxic units (TUs) for an effluent and one suspect toxicant
in a POTW effluent when two toxicants are the cause of toxicity 2-2
2-3. Correlation of toxic units (TUs) for an effluent and two toxicants in a
POTW effluent 2-3
2-4. Correct (top) and incorrect (bottom) plots of toxic units (TUs) for non-
additive toxicants 2-4
2-5. Correlation of toxic units (TUs) for a POTW effluent and the suspect
toxicant, nickel 2-5
VI
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Acknowledgments
This document presents additional information acquired since the document
entitled Methods for Aquatic Toxicity Identification Evaluations: Phase III Toxicity
Confirmation Procedures (EPA-60013-88-036; EPA, 1989D) was prepared by Donald
Mount and published in 1989. This manual reflects new information, techniques, and
suggestions made since the Phase III confirmation methods for acute toxicity were
developed. The suggestions, techniques and cautions contained in this document are
based on a large database generated by the staff of the National Effluent Toxicity
Assessment Center (NETAC) at the U.S. Environmental Protection Agency (EPA),
Environmental Research Laboratory, Duluth (ERL-D), MN. NETAC staff that provided
technical support consisted of Penny Juenemann and Shaneen Schmidt (ERL-D
staff), Joe Amato, Lara Anderson, Steve Baker, Tim Dawson, Nola Englehorn, Doug
Jensen, Correne Jenson, Jim Jenson, Elizabeth Makynen, Phil Monson, Greg
Peterson, and Jo Thompson (contract staff). Their collective experience has made
this document possible and the contributions are gratefully acknowledged. The
support through EPA's Office of Research and Development (ORD) and Office of
Water made this research possible at ERL-D.
Vii
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Section i
Introduction
The final confirmation phase of a toxicity identifi-
cation evaluation (TIE) consists of a group of steps
intended to confirm that the suspect cause(s) of toxicity is
correctly identified and that all the toxicity is accounted
for. Typically this confirmation step follows experiments
from the toxicity characterization step (Phase I) and
analysis and additional experiments conducted in toxicity
identification (Phase II) (EPA, 1991A; EPA, 1992; EPA,
1993A). However, there often may be no identifiable
boundary between phases. In fact, all three phases might
be underway concurrently with each effluent sample and
depending on the results of Phase I characterization, the
Phase II identification, and Phase III confirmation activi-
ties might begin with the first sample evaluated. Phase III
confirmation procedures should also follow after toxi-
cants have been identified by other means or when
treatability approaches are used. Rarely does one step or
one test conclusively prove the cause of toxicity in Phase
III. Rather, all practical approaches are used to provide
the weight of evidence that the cause of toxicity has been
identified. The various approaches that are often useful
in providing that weight of evidence consist of correlation,
observation of symptoms, relative species sensitivity,
spiking, mass balance estimates and various adjust-
ments of water quality.
The approaches described in this document have
been useful in TIEs at ERL-D. While the guidance pro-
vided in this manual is based largely on experience with
wastewater effluents, in general the methods discussed
are applicable to ambient waters (Norberg-King et al.,
1991) and sediment pore or elutriate water samples as
well (EPA, 1991B). However, specific modifications of
the TIE techniques might be needed (e.g., sample vol-
ume) when evaluating these other types of samples.
Confirmation is important to provide data to prove
that the suspect toxicant(s) is the cause of toxicity in a
series of samples and to assure that all other toxicants
are identified that might occur in any sample over time.
There may be a tendency to assume that toxicity is
always caused by the same constituents, and if this
assumption carries over into the data interpretation but
the assumption is false, erroneous conclusions might be
reached. That is why the correlation step (Section 2) is
accompanied by other approaches (i.e., Sections 3-9)
because each approach aids in revealing any changes in
the toxicant(s) in the confirmation phase of the TIE.
Seasonal trends in toxicants have been observed
in publicly owned treatment works (POTW) effluents and
some sediment samples. For example, organophosphate
pesticides have been observed to increase in concentra-
tions in wastewaters during the late winter and spring
months (Norberg-King et al., 1989). Therefore, the confir-
mation steps of Phase III might need to include seasonal
samples. This effort cannot always be pre-determined.
The presence of a different toxicant(s) must be consid-
ered throughout the TIE, and when samples are collected
over several months the seasonality of a suspect toxicant
should be carefully considered and studied. When reme-
dial action requires treatment changes, one must be
certain that toxicity from specific toxicant(s) is consistently
present and that the suspect toxicant(s) accounts for all
the toxicity. Treatment modifications will not necessarily
result in removal of all toxicants to acceptable concentra-
tions. If toxicity is caused by a variety of toxicants present
at varying intervals, the remedial actions that are practical
might differ from the remedial action required when toxic-
ity is caused by the same constituents consistently.
TIEs conducted at ERL-D have shown that toxi-
cants often are not additive or toxicants are present in
ratios such that the toxicity contribution by one might be
diluted out in the range of the effluent effect concentration
(e.g., LC50 orlCp value). Thus, the toxicant present at
lower yet toxic concentrations may not be readily dis-
cerned. The frequency of occurrence and impact on data
interpretation of either of the above cases was not ad-
dressed previously (EPA, 1989D) but are now discussed
in Section 2. Toxicants that do not express their toxicity
because of the presence of other toxicants (either the
toxicants are non-additive or the toxicants occur in dispar-
ate ratios) are referred to as hidden toxicants (Section 9).
Detection of hidden toxicants is one of the most difficult
aspects of confirmation. It is a mistake to search for a
concentration of any chemical present in the effluent at a
toxic concentration and to declare any found as the cause
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of toxicity. Matrix effects of the effluent samples make
conclusions such as these subject to error without further
work as either the hidden toxicant(s) or the principal
toxicant(s) are likely to be missed using such an ap-
proach.
There is a strong tendency to shorten or eliminate
the confirmation steps because by the time Phase III
confirmation has been reached, the investigators might
be convinced of the cause of toxicity and the confirmation
steps seem redundant. However, one cannot expect to
concentrate the effluent on a C18 solid phase extraction
(SPE) column and not change a complex mixture such as
effluents, and arrive at some false conclusions about the
toxicants in the earlier phases.
Not all approaches discussed in the following
sections will be applicable to every effluent, and addi-
tional approaches might need to be developed during the
TIE. The various approaches need not be performed in
any particular sequence, and the list of possible ap-
proaches will get larger as experience is gained. To
effectively evaluate effluent samples from one particular
discharger to obtain a correlation, substantial calendar
time could be required and any steps for correlation
should be initiated at the beginning stages of Phase III.
Judgement must be made as to how many of the ap-
proaches described in Phase III confirmation should be
used and how many samples for each shauld he oarrir
pleted. How completely Phase III confirmation is done will
determine the authenticity of the outcome. The amount of
confidence in the results of the TIE that is required is
dependent at least in part on the significance of the
decision that will be based on the results. For example, if
a suspect toxicant can be removed by pretreatment or by
a process substitution, a higher degree of uncertainty
may be acceptable than if an expensive treatment plant is
to be built. Such considerations are subjective and cannot
be reduced to a single recommended decision making
process with a specified number of samples.
Time and resources might be conserved if identi-
fication (Phase II) and confirmation (Phase III) can be
started on the very first effluent sample used in the Phase
I characterization. However, this is only possible when the
results from the Phase I characterization are definitive
enough to allow the investigators to proceed to identifica-
tion and confirmation. In the acute Phase III confirmation
document (EPA, 1989D), although perhaps not explicitly
stated, performing Phase I characterizations on several
samples before attempting Phases II and III was implied.
Initiating the Phase III confirmation steps earlier in the TIE
is often particularly useful. In addition, many regulatory
agencies have adopted a policy that requires that the
previous TIE approach be modified. For some discharg-
ers, action might be required after the first exceedence in
toxicity, which means that each effluent sample collected
for toxicity testing is of equal regulatory concern when the
toxicity is greater than the permit allows. This regulatory
practice was not in place in 1989 when the earlier TIE
guidance was available (EPA, 1989D) and at that time we
did not expect that the cause of toxicity in one sample
could be sufficiently deduced as we have been able to do.
The importance of confirmation on several samples is not
reduced by the importance of conducting confirmation
steps on single samples; rather, the cause of toxicity for
each sample must be confirmed.
In addition to the importance of each sample with
toxicity greater than the allowable amount specified in a
permit, a sample that is quite different from the previous
samples must be evaluated to determine if the data point
must be included in the Phase III correlation final data
analyses. For each effluent sample, the data points must
be explainable. If one sample is quite different than other
samples it can cause the correlation to be less useful;
however, if it can be shown to have a different toxicant the
data point for that sample can be eliminated from the
correlation. For example, suppose five consecutive
samples during a Phase III evaluation exhibited toxicity
that correlated well with a suspect toxicant. Then a sixth
sample exhibits greater toxicity than previous samples
while the measured concentration of the suspect toxicant
is much lower than measurements on previous samples.
In this sixth sample, the greater toxicity is thought to be
caused by a different toxicant. Now in plotting the data for
the correlation (Section 2), the datum point for the sixth
^am^a wJJ, •?<& t/& tjmilati to 'frfe pOTte for *rrie existing
regression and could render the correlation non-signifi-
cant. If however, when the sixth sample is then subjected
to intensive study using Phase I characterization and
Phase II identification techniques, and if another toxicant
is identified (or even if Phase I only shows that the toxicity
has very different characteristics), datum for the sixth
sample can legitimately be excluded from the correlation.
This preserves the worth of the data for the previous five
samples. In confirmation, every effort should be made to
determine why a particular sample shows different re-
sponses in the various TIE steps from other samples.
This is not to imply that multiple effluent samples
need not be subjected to Phase I manipulations, even if
Phase II and/or Phase III are initiated on the first sample.
Most effluent samples tend to be representative of the
routine effluent discharge. However, determining what is
the characteristic discharge for each effluent is important
to the final success and completeness of the TIE.
When Phase III is completed, all results that were
obtained during the TIE should be explainable. Unless the
results make sense for all samples (aside from an occa-
sional aberrant data point) something has been missed or
is wrong. If so, the confirmation is not complete. Many
techniques used in Phase III require keen observations
and extensive or broad knowledge of both chemistry and
toxicology but above all the ability to synthesize small bits
of evidence in a logical sequence is essential. This TIE
work is most effective when scientists interact daily.
1-2
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A note of caution. If data obtained on early samples
during Phase I are to be used for Phase III purposes,
quality control will have to be suitable to provide defen-
sible data (cf., EPA, 1991A; EPA, 1992; EPA, 1993A). In
Phases I and II, the permissibility of using small numbers
of animals and replicates, and omitting measurements
such as pH, DO, and temperature that are required for
routine monitoring tests or single chemical tests was
discussed (EPA, 1989E; EPA, 1991 A; EPA, 1992; EPA,
1993A). These modifications were made to reduce cost
and allow more testing, but at this point shortcuts must be
avoided because definitive data that constitute the basis
for important decisions are generated in Phase III. For
Phase III testing, the effluent test protocols that triggered
the TIE (EPA, 1991C; EPA, 1993B) should be followed,
paying careful attention to test conditions, replicates,
quality of test animals, representativeness of the effluent
samples tested, and strict QA/QC analytical procedures
including blanks and recovery measurements. Analytical
work must be selective for the identity of the toxicant and
its concentration measurement. When small differences
in toxicity must be detected, concentration intervals should
be smaller to obtain partial effects (e.g., use dilution
factors of 0.60 or 0.65 versus 0.5). Remember, all of the
data from Phases I and II (for either acute or chronic
toxicity) are considered preliminary relative to Phase III
data. However, if a suspect toxicant is identified and
Phases I and II data may be necessary for confirmation,
stricter QA/QC can be applied for each of the subsequent
Phases I and II techniques so that the data can be used in
Phase III.
For samples exhibiting chronic toxicity, modifica-
tions or 'changes to some of the TIE procedures are
required for confirming the cause of chronic toxicity. Re-
member that for confirmation (as well as for Phases I and
II), only a single sample of effluent should be used for
each renewal in any chronic test (cf., EPA, 1992; EPA,
1993A). This is important because one cannot correlate a
measured concentration of a toxicant with the toxicity
measured in a test if multiple samples are used for each
renewal and the toxicant is not present in some samples
but other toxicants appear. Even more likely, the ratios of
the toxicants, when more than one is present, might
change from sample to sample. In these instances, there
is no valid way to calculate the toxicity of a given toxicant.
Overall, considerations for chronic toxicity tests in Phase
III are not much different than acute toxicity tests in Phase
III. At present, permit requirements specify the 7-d test
and unless data are gathered to show that the 4-d and 7-
d tests yield the same results and that the same toxicants
are involved, the 7-d test should be used for confirmation
(cf., EPA, 1993A). If the 4-d Ceriodaphnia dubia test has
been used instead of the 7-d C. dubia test (see EPA,
1992) during Phases I and II, serious consideration should
be given to returning to the 7-d test for Phase III.
When identification of the toxicant(s) causing
chronic toxicity is desired, and the effluent also exhibits
acute toxicity, it might be possible to use acute toxicity as
a surrogate measure to characterize the toxicity in Phase
I and assist in an identification in Phase II. It must be
demonstrated that the cause of the acute toxicity is the
same toxicant(s) as the toxicant(s) causing the chronic
toxicity. Yet for confirmation, use of chronic toxicity end-
points to confirm the cause of the chronic toxicity is
strongly recommended to avoid misleading the TIE re-
sults when using acute toxicity as a surrogate for chronic
toxicity. As discussed in the chronic Phase I manual
(Section 5.8; EPA, 1992), effect levels for chronic tests
should be calculated using the linear interpolation method
rather than the hypothesis test (EPA, 1992). In order to
get more precise estimates of endpoints, test concentra-
tion intervals might have to be narrowed (see above).
However, when point estimation techniques for other than
survival endpoints (such as the inhibition concentration
(ICp); EPA, 1993B) are used, a point estimate effect
concentration can be estimated. The effect concentration
estimates will also be more accurate when intermediate
concentrations are used (i.e., use dilution factors of 0.6 or
0.65).
1-3
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Section 2
Correlation Approach
2.1
Correlation
The purpose of the correlation approach is to
show whether or not there is a consistent relationship
between the concentration of suspect toxicant(s) and
effluent toxicity. For the correlation approach to be useful,
the toxicity test results with the effluent must demonstrate
a wide range of toxicity with several effluent samples to
provide an adequate range of effect concentrations for
the regression analysis. For sediment samples, spatial
variability might be used to perform correlation analyses
(EPA, 19916).
The effluent effect concentration (i.e., LC50 or
ICp) data and the measured toxicant concentration data
must be transformed to toxic units (TUs) for the regres-
sion analysis to evaluate whether or not a linear relation-
ship exists. Effluent TUs are obtained by dividing 100%
by the effect concentration expressed in percent of the
effluent (cf., EPA, 1991A; EPA, 1992). The suspect toxi-
cant concentration is converted to TUs by dividing the
measured toxicant concentration by the LC50 or ICp for
that toxicant (data to make this comparison might have to
be generated; EPA, 1993A). If more than one toxicant is
present, the concentration of each one is divided by the
respective LC50 or ICp value and the TUs can then be
summed (cf., discussion below for non-additive toxicants).
Most of the effluents we have tested have exhib-
ited a wide range of toxicity with several different samples
and therefore the data can be used in the correlation
approach. Typically for the correlations that we have
conducted, the data used are from toxicity tests without
any manipulations and from chemical measurements on
the effluent samples for the concentrations of the suspect
toxicant. However for effluents where ammonia was the
cause of the toxicity, the effluent toxicity. results have not
varied in toxicity enough, nor have the ammonia concen-
trations fluctuated enough to use the data in a correlation.
Also, when the effect concentration is greater than 1 00%,
this information is not useful since the data point cannot
be included in the regression analysis. However, when
samples are marginally toxic or when the suspect toxicant
concentrations do not vary enough from sample to sample
(i.e., ammonia is cause of toxicity), changes in toxicity can
be induced by sample manipulation (cf., EPA, 1993A) and
this toxicity data can be used to develop a different type of
correlation. For example, the toxicity of a given amount of
total ammonia can be changed by over an order of
magnitude by altering the pH of aliquots of the effluent
within an acceptable physiological range (e.g, pH 6 to 9).
For some metals and some species, the toxicity can also
be changed by adjusting the pH and using dilution waters
of varying hardness. This type of data is useful in the
correlation step as providing additional weight of evi-
dence. Therefore, the idea of minimal manipulation(s)
and any risk of creating artifactual toxicity are off set by the
utility of the data.
An example of the regression from am effluent
from a POTW in which the suspect toxicant was diazinon
is given in Figure 2-I. The independent variable (x-axis) is
the TUs of diazinon and the dependent variable (y-axis) is
the effluent TUs. The solid line is the observed regression
line obtained from the data points, and the dashed line is
the expected or theoretical regression line. If there is 1 .0
TU of the toxicant in 100% effluent, then the effluent
should have 1 .0 TU (i.e., the LC50=100%). Likewise for
2.0 TUs of suspect toxicant, the effluent TUs should be
2.0, et cetera. Thus, the expected line has a slope of one
and an intercept of zero. In Figure 2-1, the intercept (0.19)
is not significantly different from zero and the slope is very
close to 1 (1.05). The r2 value is 0.63 which, while not
high, indicates that the majority of the effluent toxicity is
explained by the concentration of the toxicant. As the r2
becomes lower, less confidence can be placed on slope
and intercept. In a small data set such as this, one datum
point that had 5.0 TUs for the effluent toxicity lowered the
r2 value substantially. As discussed in Section 1, if an
intensive effort had been expended on that sixth sample
and another toxicant(s) had been found, this particular
datum point could have been excluded and the r2 value
would have been higher.
In another POTW effluent, diazinon was also the
suspect toxicant. For these data (Figure 2-2), the slope is
1.38, the intercept is 1.24 and the r2 value is only 0.15,
which all indicate poor fit for diazinon as the only toxicant.
The low r2 value indicates a large amount of scatter,
therefore little can be inferred from the slope and the
intercept. Based on this correlation, we returned to Phase
II analytical procedures and identified two other organo-
phosphates (chlorfenvinphos (CVP) and malathion). Tox-
icity data indicated that CVP was present at toxic
2-I
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LLJ 3
"5
r2 = 0.63
slope = 1.05
y-intercept = 0.19
0123456
TUs of Suspect Toxicant(s)
Figure 2-1. Correlation of toxic units (TUs) for an effluent and one
suspect toxicant in POTW effluent.
slope = 1.38
y-intercept = 1-24
Z 3 4 5 6
TUs of Suspect Toxicant(s)
Figure 2-2. Correlation of tox'c units (TUs) for an effluent and cne
suspect toxicanr in a POTW effluent when two toxicants
are the cause of toxicity.
concentrations while malathion was no?. After testing
each compound both separately and as a mixture, the
toxicity from all three chemicals was determined to be
additive, so a new correlation was begun with analytical
measurements made for all three chemicals. CVP and
diazinon have nearly identical LC50 values for the spe-
cies (C. dubia) used in this TIE. Malathion is about one-
fourth as toxic as CVP or diazinon. Since the measured
concentrations of malathion were lower than its toxicity, it
was not included in the regression analysis. In a new
correlation with data for the TUs summed for CVP and
diazinon versus the effluent TUs, the data show a much
better fit to the expected slope and intercept and a high r2
value (Figure Z-3). Malathion TUs could also have been
included in the regression (although its contribution to
toxicity was minimal) because it was additive with other
toxicants. This type of situation is discussed below.
In addition to slope and intercept, some judge-
ment of the scatter about the regression line must be
made. This can be done statistically, but when the sample
size is large, the scatter can be very large and yet not
negate the relationship. A-suggested approach to avoid
the effect of sample size on the significance of scatter is
to set a lower limit on r2. This value (often expressed as
percent) provides the measure of how much of the ob-
served effluent toxicity is correlated to the measured
toxicant. It is not dependent on choosing the correct effect
concentration of the toxicant. The specific choice of the
minimum value of r2 should be made based upon the
consequences of the decision. It is important to recognize
that experimental error makes an r2 value greater thgn
0.80 or 0.85 difficult to obtain. Therefore, where minimal
chance of an incorrect decision is required, an r2 value of
nearly 0.80 may be used. Where an increased risk of an
incorrect decision (i.e., a lesser amount of the toxicity
accounted for) is acceptable, a lower value such as 0.60
may be used.
Since <1 .0 TU cannot be directly measured in the
effluent, such values are, of necessity, excluded from the
regression. (This comment is exclusive of the use of
concentrates such as the C13 SPE fractions' where TUs of
<1.0 are possible.} However in some instances, when the
TUs based on chemical analyses are <1.0 TU and efflu-
ent effect values are <1.0 TU, the data support the validity
of the regression provided a suspect toxicant has been
found in several previous samples. In the correlation for
the effluent toxicity depicted in Figure 2-2, toxicity was
present in a different fraction (Phase II non-polar organic
identification) than where the pesticides were identified. A
specific toxicant was not identified in that fraction and
toxicity was not always measurable in that fraction. How-
ever, this additional toxicity may have decreased the r2
value.
Correlation might be more definitive when two or
more toxicants are present. For example, suppose three
toxicants are involved. If each toxicant has the same
LC50 and each is strictly additive with the ratio of their
concentrations remaining the same, the slope will be the
expected but the intercept will be positive if all toxicants
TUs can be calculated from toxicity tests with the fractions, the
concentrate or the HPLC fractions as described in Phase II (EPA,'
1993A).
2-2
-------
o
0)
f2=0.73
slope = 0.82
y-intercept = 0.46
0 1.2 3 4 5 6
TUs of Suspect Toxicant(s)
Figure 2-3. Correlation of toxic units (TUs) for an effluent and two
toxicants in a POTW effluent.
are not identified. If the relative amounts (ratios) of each
toxicant vary from sample to sample, the slope, intercept
and r2 will be different from the expected if only one
toxicant is identified. If the toxicity of one of the toxicants
is substantially different, and if the ratios of the three
toxicants vary from sample to sample, then the slope,
intercept, and r2 value will all be different from expected if
all are not identified. Much can be learned from studying
the interrelationship of slope, intercept and the r2 value.
For example, a high r2 value and an intercept near zero
with a slope larger than 1 can be caused by using an
effect concentration for the toxicant that is not appropriate
for the toxicant in the effluent matrix (e.g., suspect toxi-
cant is more toxic in effluent matrix than in single chemi-
cal test). This error causes the toxicant TUs to be too few
relative to the effluent TUs (Figure 2-4) (cf., discussion
below on non-additive toxicants). If toxicant concentra-
tions and effluent toxicity show a wide distribution, a
significant correlation will be easier to demonstrate than
for a narrow range.
Great care must be taken to understand whether
or not toxicants are additive or if the TUs for each toxicant
are so different that only one toxicant determines the
effect level. For either situation, the resulting data will
have to be interpreted as though the toxicants are non-
additive. For example, suppose the ratio of TUs is so
disparate that at the effluent effect concentration, the
toxicant with fewer TUs is always present at a fraction of a
TU (e.g., 0.25 of a TU). Whether the two toxicants are
additive or not is irrelevant because the major toxicant will
set the effluent effect concentration. While 0.25 TUs of
the minor toxicant appear to be relatively unimportant in
view of experimental variability, this affects the regres-
sion. If in one sample the effect concentration is 25% and
the 4 to 1 ratio of toxicants occurs, there are 4 TUs of the
major toxicant and 1 TU of the minor toxicant. If the
toxicant concentrations are summed, 5 TUs will be plotted
against 4 effluent TUs, and this results in a 25% error.
When secondary toxicants are present in concentrations
that will not contribute to the effect concentration of the
effluent, they should not be included in the correlation
data set. Obviously if an effluent had several toxicants in
dissimilar ratios, the error of including the minor TUs in a
correlation plot could be large and may negate the corre-
lation significance. The investigator should evaluate the
data in regression plots to consider the significance of the
contribution of the secondary toxicant especially if the
toxicants appear to be additive.
Unfortunately the minimum fraction of a TU that is
detectable will depend on the precision of the laboratory
performing the testing. And of course the precision of the
testing is not only dependent on the quality of the work,
but the inherent precision of measuring specific toxicant
TUs. That is, the toxicity measurement for some chemi-
cals is more precise than for some other chemicals. In
general, a chemical such as NaCI whose toxicity is gener-
ally not affected by pH, alkalinity, hardness, total organic
carbon (TOC), suspended solids or solubility, can be
measured more precisely than a chemical whose toxicity
is affected by these factors, such as lead or copper.
Therefore, each laboratory must determine which frac-
tional value of a TU at the effect concentration is
unmeasurable, thus indicating which TUs contributed by
the minor toxicant should be deleted from the correlation
data set.
Clearly, if two or more toxicants are strictly non-
additive, then only the major one (the one present in the
most TUs) should be included in the correlation data set.
Since additivity might be easier to measure than the
minimum measurable contribution of a fraction of a TU, it
may be preferable to first determine if additivity occurs. If
substances appear to be partially additive, then very
careful work is required to properly add TUs.
Some very unusual decisions are required in
accepting data into the correlation database when toxi-
cants are strictly non-additive. For example, consider zinc
and ammonia in the same effluent sample; we have fcund
them to be strictly non-additive. Also consider that in
some samples zinc and ammonia occur in TU ratios of 3
to 1 and in other samples the ratio is 1 to 2. In the
regression for the 3 to 1 ratio samples, only zinc TUs
should be plotted. In the regression for the 1 to 2 ratio
samples, only ammonia TUs should be plotted. For this
particular example, 3 TUs for the first sample and 2 TUs
for the second sample would be used if the data is
interpreted correctly (i.e., plotting total TUs) or 4 and 3
TUs would be used respectively, if the data is interpreted
2-3
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c
0>
3
Correct Plot
I I I I
1234
Calculated TUs
4 -
1 —
Incorrect Plot
\ \
I
2 3
Calculated TUs
Figure 2-4. Correct (top) and incorrect (bottom) plots of toxic units
(TUs) for non-additive toxicants.
incorrectly. The slopes for both plots would be 1 but a
negative intercept instead of an intercept of 0 would be
obtained for the incorrect plot. The more similar the TUs
of each toxicant are to each other, the greater the error in
the correlation will be.
2.2 Correlation Problems Caused by Matrix
Effects
Correlation becomes much more difficult when
the toxicants interact with the other effluent constituents
in ways that change their toxicity and we refer to these
changes as matrix effects. There are numerous matrix
effects and all of them will not be discussed here; instead
a framework is provided to aid in designing tests or test
conditions to validly incorporate matrix effects in such a
manner that useable correlation data can be obtained.
Matrix effects generally fit into one of two catego-
ries. One category is when the toxicants change form in
some manner which exhibit a different toxicity. A very
common example is ammonia which changes from NH, to
NH,+ as pH decreases. NH,+ is so much less toxic than
NH, that it is often considered nontoxic? Another example
is HCN whose most toxic form is as un-dissociated HCN,
a form predominating at low pH values. As pH increases
the equilibrium shifts to more H* and CN-. If metals are
present, metal-cyanide complexes form which are often
less toxic than HCN but metal-cyanide complexes might
vary in toxicity depending on the metal. For example, iron-
cyanide complexes are much less toxic than some of the
other metal complexes. Metal-cyanide complexes might
also photodecompose in sunlight releasing HCN or H*
and CN-, depending on pH.
A second category of matrix effects involves such
physical changes as sorption or binding in some manner
so as to make the toxicant unavailable to the organism.
For example, non-polar organics sorb onto suspended
solids, and some metals, such as copper, also sorb onto
suspended solids. The presence of organic matter on
suspended solids might increase the sorbtive capacity.
Predictably, changes in water chemistry often change the
sorption/solution equilibrium and thereby, change the por-
tion of total toxicant that is available to the organism.
To further complicate matters, biological charac-
teristics of the test organisms might change the availabil-
ity of the same toxicant form. For example a non-polar
organic sorbed on suspended solids such as bacterial
cells, might be unavailable to a fish but readily available to
daphnids because cells might be ingested and digested
by daphnids. The uptake route then is through the diges-
tive tract but the toxicant has entered the body none-the-
less.
From the above discussion, it is obvious that one
method of correlation will not be applicable for all toxi-
cants. A temptation may be to remove the toxicant from
the effluent and then use the effluent as a diluent to
measure toxicity. However, because effluents are so com-
plex and undefined, there is virtually no way to remove
one or a few constituents and still be certain other charac-
teristics have not been changed. For example, zeolite
removes ammonia but it also removes some metals and
non-polar organics; the C1fl resin removes metals as well
as non-polar organics; ion exchange columns remove
ionized constituents, but non-polar organics also are re-
tained by the columns. Toxicant removal procedures have
utility but require very complicated simultaneous testing
of the effluent and proper blanks (cf., EPA, 1992; EPA,
2See specific discussion in Section 3, Phase II (EPA, 1993A).
2-4
-------
1993A) is necessary to properly interpret results (cf.,
Section 9 on hidden toxicants).
In Phase III, quantitative comparisons are being
made between toxicity and concentrations of toxicants
rather than qualitative comparisons as in Phases I and II
(EPA, 1991A; EPA, 1992; EPA, 1993A). In the correlation
approach, such comparisons are the essence of the
technique. Therefore even small changes in form or avail-
ability might be unacceptable. This means that manipula-
tions and changes must be minimized when effluent
toxicity and toxicant concentrations are to be compared.
Solvent extraction, so commonly used for organic
analyses, is likely to extract biologically unavailable or-
ganics as well as soluble forms. The total measured
concentration may be larger than the true exposure con-
centration. Use of the C SPE column also is not free
from problems as the C18 SPE column is a finer filter than
the glass fiber filters commonly used for pre-column
filtration. Therefore solids are likely to be physically re-
tained on the upper part of the column. When the column
is eluted with methanol, the methanol extracts toxicant(s)
from the solids (which might not be biologically available)
as well as elutes the C18 sorbent itself. For Phases I and
II, this might be unimportant, but for the Phase III correla-
tion step where careful quantitative comparison is neces-
sary, the effect might be unacceptable. Such problems
probably reach a maximum when working with samples
such as highly organic sediment pore water (with high
organic characteristics) where much of the chemical might
be biologically unavailable.
The central problem for either type of matrix
effect is the difficulty of analytically measuring the biologi-
cally available portion of the specific toxic form. A correla-
tion for a POTW effluent where for nickel was suspected
of causing the toxicity is shown in Figure 2-5. During
Phase I, the acute toxicity was removed with EDTA
additions, and in Phase II the nickel was measured at
toxic concentrations to C. dubia. The toxicity correlated
very well with total nickel concentration (r2 = 0.89 and a
slope of 1 .17) and it appeared that only nickel seems to
be involved. But the intercept of -12.34 is quite different
from the expected zero. Such an intercept would be
expected if there were a relatively fixed amount of nickel
which was not biologically available in all samples. In this
example, because all other confirmation data corrobo-
rated nickel as the toxicant, a constant concentration of
nontoxic nickel was thought to provide the explanation for
the unexpected intercept value. However, there is no
obvious reason to think that the quantity, or even the
percentage of total toxicant, is the same across samples
for other toxicants, or for nickel in other matrices.
For the effluent samples that lose their toxicity in
a short time, the nontoxic effluent can be used for the
suspect toxicant(s) tests as a diluent in parallel tests
using a standard dilution water to elucidate matrix effects
on toxicity. Toxicity test results with quite different toxicity
would reflect matrix effects. If toxicity is persistent, devel-
50
40
30
_
UJ
20
10
Observed
Theoretical
r2 = 0.89
slope = 1.17
y-intercept = -12.34
10 20 30 40
TUs of Suspect Toxicant (Nickel)
50
Figure 2-5. Correlation of toxic units (TUs) for a POTW effluent and
the suspect toxicant, nickel.
oping two separate correlations using pure chepiical addi-
tions on two different effluent samples, each with sub-
stantially different toxicant concentrations, might be useful.
If the toxicity test results indicate that the biologically
unavailable portion changes with measured concentra-
tions, the slope should be different than one. This ap-
proach requires careful work and the investigator must
consider incorporating equilibrium time experiments (cf.,
EPA, 1993 A).
Metals can be especially difficult toxicants to
implicate using correlation because the toxicity of metals
is typically very matrix dependent. When the knowledge
of these characteristics is extensive for a chemical, as it is
with ammonia (see Phase II), testing can be tailored to
the chemical and a very powerful correlation obtained.
The large amount of available information on ammonia
does not exist for most metals. In these instances, the
logic pattern should to be reversed where the approach
has to become: if x ;s the toxicant, what are the matrix
effects?. These can be found by pure chemical testing
combined with Phases I or II manipulations. Once an
adequate understanding of matrix effects is obtained, the
information can be used to answer the question: Is the
effluent toxicant behavior consistent with the matrix ef-
fects for the suspect toxicant?
Matrix effects will have varying impacts on toxi-
cant behavior that also depends on the effluent effect
concentration. For effluents which have effect concentra-
tions in the <10% range, the test solutions will more
closely resemble the diluent water matrix than the efflu-
ent. If the effluent has effect concentrations in the 50% to
100% range, the matrix effects of the test solution will
most likely resemble those of the effluent, not of the
dilution water. Since effluent TUs are calculated from
2-5
-------
responses occurring in the dilution near the effect con-
centration, the matrix characteristics of that concentration
are of the most concern for correlation. Thus the impor-
tance of the effluent matrix effects diminishes as the
toxicity of the effluent is greater (i.e., matrix at effect level
is more like dilution water).
One can safely say that the difficulty of simulating
the matrix effects with a simulated effluent is quite large
so that the choice is clearly to use the actual effluent
when possible. An important reason for this choice is that
so few matrix effects have been studied extensively, and
beyond pH and hardness little data exists. Even then the
interrelationship between pH, alkalinity and hardness were
often ignored.
The above discussion does not provide all of the
options on how to handle matrix effects. However, it
should provide convincing evidence that more than the
correlation step alone is necessary to provide adequate
confirmation!
In summary, the TIE research experience has
revealed two major areas of potential problems in using
the correlation approach. The lack of additivity for toxi-
cants found in effluents requires careful analysis when
calculating TUs for regression purposes. Secondly, when
there are matrix effects, correlation becomes difficult be-
cause the effluent matrix might change from sample to
sample and because there are no analyses specific for
the toxic forms. For such effluents, other confirmation
techniques should be used more extensively to better
support the overall confirmatory efforts.
2-6
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Section 3
Symptom Approach
Different chemicals may produce similar or very
different symptoms in a test species. Probably no symp-
tom of intoxication is unique to only one chemical. There-
fore, while similar symptoms observed between two
samples means the toxicant(s) could be the same or
different, different symptoms means the toxicant(s) is
definitely different, or there are multiple toxicants in the
two samples. By observing the symptoms displayed by
the test organisms in the effluent and comparing them to
the symptoms displayed by test organisms exposed to
the suspect toxicants, failure to display the same symp-
toms means the suspect toxicant(s) is probably not the
true one or the only one.
Behavior of most test species is difficult to put
into words so that a clear image of behavior is obtained.
Behavioral and morphological changes of 30-d old fathead
minnows (Pimephales promelas) were used as diagnos-
tic endpoints in 96 h flow-through single chemical tests.
Organic chemicals of various modes of action were tested
and video recordings were used to monitor the behav-
ioral response (Drummond et al., 1986; Drummond and
Russom, 1990). Substances within a single chemical
classification did not necessarily cause the same type of
response (Drummond and Russom, 1990). Therefore, it
is difficult to predict chemical classification using behav-
ioral monitoring alone.
This type of behavioral monitoring data does not
exist for the cladocerans or the newly hatched fathead
minnows or other species that are most frequently used
in the TIE process. However, noting various symptoms is
useful in the TIE. This is done by simply exposing the test
species to the suspect toxicant(s) and observing how
they react. By the time confirmation is initiated, toxicity
tests with the suspect toxicants will have been conducted
using pure compounds and symptoms may have been
observed. It is important to note the symptoms observed
during all testing because such characteristics can be
very helpful in confirmatory work.
The intensity of exposure concentrations might
change the symptoms observed with the suspect toxicant
in the effluent. Therefore, it is important to compare
symptoms at concentrations that require about the same
period of onset. This can be done by comparing symp-
toms at exposure concentrations that have similar TUs. In
this way both the unknown (sample) and the known
toxicants (pure compound) can be set at the same toxicity
level.
Observations of the organisms should not be
delayed until the normal length of the test has elapsed.
With some toxicants, the test organisms will show distinc-
tive symptoms soon after the exposure begins, whereas
later, symptoms are often more generalized and less
helpful. For some other toxicants, a sequence of different
symptom types are displayed by the test organism over
the exposure period and the sequence may be more
definitive for a given chemical than the individual symp-
toms. In few cases will the symptoms be unique enough
to specifically identify the toxicant, but symptoms different
from those caused by the pure suspect toxicant are
convincing evidence that the suspect toxicant is not the
true or only one.
A second caution is needed regarding mixtures of
toxicants. Mixtures of toxicants can produce symptoms in
test animals different from the symptoms of the individual
toxicants comprising the mixture. When more than one
toxicant is involved, the investigator must not only include
all the toxicants, but include them in the same ratio as
measured in the effluent. Often the toxicant of the mixture
at the highest concentration relative to its effect concen-
tration will cause most of the symptoms. As for single
toxicants, the mixture concentration causing the same
endpoint in a similar exposure period should be com-
pared. Spiking effluent with the suspect toxicants and
comparing the results of the spiked effluent sample and
the unspiked effluent sample toxicity tests, both near their
effect concentrations, is a good approach to take (Sec-
tion 5).
Symptoms caused by the toxicant(s) might be
quite different among different species of organisms;
therefore the use of two or more species provides in-
creased defmitiveness of the observations. For both spe-
cies, the researcher must compare symptoms at
concentrations that are equitoxic. The greater the differ-
ence in sensitivity, the more important this becomes, The
chemical concentration is unimportant; the important con-
sideration is that equitoxic concentrations are compared.
3-I
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Suppose, for example, species A and B have LC50
values for a suspect toxicant of 1 and 80 mg/l. Then
concentrations of 2 and 160 mg/l may be used to com-
pare symptoms of species A and B, respectively. If the
onset of symptoms is rapid, then perhaps 1.25 and 100
mg/l(1.25xLC50) should be tried. Since symptoms vary
with the exposure intensity, using various multiples of the
LC50 (i.e., 0.5, 1, 2x) can add additional confirmation
data, if the same set of symptoms are seen in both series.
If more than one toxicant is involved, and the ratio of the
two species' LC50 values for toxicant A is markedly
different than for toxicant B, C, D, . . . . then the definitive-
ness of using symptoms is even greater.
For acute toxicity, time-to-mortality at equitoxic
concentrations can be used as a symptom type of test.
Some chemicals cause mortality quickly and some cause
mortality slowly. If for two effluent samples, toxicity is
expressed quickly for one and for the other very slowly,
the toxicants are probably not the same.
In chronic testing, use of symptoms is also appli-
cable. For example, adult mortality, number of young/
female, death of young at birth, growth retardation, abor-
tion, or time to onset of symptoms, all can also be
monitored and such observations may be useful. The
shape of the dose response curve may also be a determi-
nant in assisting in confirmation. Some chemicals show
an all or none type of response (diazinon) while others
(i.e.. NaCI) display a relatively flat concentration-response
slope for chronic toxicity.
3-2
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Section 4
Species Sensitivity Approach
The effect concentrations can be compared for
the effluent of concern and the suspect toxicants, using
species of different sensitivities. If the suspect toxicant(s)
is the true one(s), the effect levels of effluent samples
with different toxicity to one species will have the same
ratio as for a second species of different sensitivity. Also
the ratio for each species should be the same as for
known concentrations of the pure toxicant. The same
ratio of effect values for two species implies the same
toxicant in both samples of effluent. Obtaining the same
effluent toxicity ratio among various effluent samples for
each species as is obtained by exposure to comparable
concentrations of known toxicants, implies that the sus-
pect toxicants are the actual ones present. However, if
other effluent characteristics affect toxicity and if they
vary, the ratios could also be affected.
The common notion that goldfish are resistant to
most toxicants and trout are sensitive to most toxicants is
not readily substantiated (AQUIRE, 1992). Many species
are more sensitive to certain groups of toxicants than
trout. Of course, there are generalizations that can be
made. For example, sunfish (Centrarchids), frequently
are much more resistant to metals than goldfish, min-
nows, and daphnids (AQUIRE, 1992). Daphnids tend to
be more resistant to chlorinated hydrocarbon insecticides
than many fish species and more sensitive to organo-
phosphate insecticides (AQUIRE, 1992). These differ-
ences must always be verified for the suspect toxicants;
generalities can only be used as an initial guide to
species selection. Sensitivity differences of 1 0-i OOx may
occur in some chemical groups and not in others. If
several toxicants are involved, interpreting the results
and designing the ancillary experiments is more difficult.
If successful, the power of the result for multiple toxicants
is much greater than for a single toxicant. The difference
in sensitivity between Ceriodaphnia and fathead min-
nows has, on several occasions, revealed either a change
in the suspect toxicants present in a series of effluent
samples, or the presence of other toxicants in addition to
those suspected.
Comparison of sensitivity among species has
another very important use. Some species may evidence
toxicity from an effluent constituent that the TIE test
species did not. If this happens, then the above compari-
son will be confused, but at least there will be a warning
that the suspect toxicant may not be the cause of toxicity.
In order to determine what is happening, the investigator
should step back to Phase II, and possibly step back to
Phase I to characterize the additional toxicant and then
identify the toxicant using the new species. A second
Phase III effort might be necessary for this toxicant and
species. It is important not to assume that th* resident
species have the same sensitivity as the TIE test species.
Especially for freshwater discharges into saltwater this
concern is critical when a saltwater organism triggered
the TIE, because at present the techniques and proce-
dures described in Phases I and II are most likely to be
done using freshwater organisms especially since the
effluent is freshwater. If the concern is for marine organ-
isms and their protection cannot be assumed (cf., Section
8, Phase I; EPA, 1991A), confirmation must be conducted
with marine organisms.
In chronic testing, chemical and physical condi-
tions might differ more among tests on different species
because food must be provided during the test period and
different foods are used for each species. For example,
the final pH of fathead minnow 7-d tests might be lower
than in acute fathead minnow tests and both are likely to
be lower than in Ceriodaphnia chronic tests due to greater
respiration rates for fish than cladocerans and food in fish
tests. If the investigation was to confirm ammonia toxicity,
this pH difference could result in confusing results by
showing the Ceriodaphnia to be more sensitive than the
fathead minnows when the reverse should be true (cf.,
EPA, 1993A; Phase II). The above example illustrates
reasons to maintain careful quality control in Phase III
work.
4-I
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Section 5
Spiking Approach
In spiking experiments, the concentration of the
suspect toxicant(s) is increased in the effluent sample
and then toxicity is measured to see whether toxicity is
increased in proportion to the increase in concentration.
While not conclusive, if toxicity increases proportionally
to an increase in concentration, considerable confidence
is gained about the true toxicant( Two principles form
the basis for this added confidence. To get a proportional
increase in toxicity from the addition of the suspect
toxicant when it is in fact not the true toxicant, both the
true and suspect toxicants would to have 1) very similar
toxicity and 2) to be strictly additive. The probability of
both of these coinciding by chance is small.
Removing the suspect toxicants from the effluent
without removing other constituents or in some way
altering the effluent is usually not possible. The inability
to do this makes the task of establishing the true toxicity
of the suspect toxicants in the effluent difficult. For many
toxicants, effluent characteristics, such as TOC, sus-
pended solids, or hardness, affect the toxicity of a given
concentration, Some characteristics, such as hardness,
can be duplicated in a dilution water, but certainly not
TOC or suspended solids because there are many types
of TOC and suspended solids, and generic measure-
ments do not distinguish among the different types. For
example, effluent TOC occurs as both dissolved and
suspended solids. In POTW effluents, the source of the
TOC is likely to be largely from biological sources, both
plant and animal (e.g., bacteria) and bacteria are likely to
make up a large component of suspended solids. If there
have been recent storms, oily materials from stormwater
runoff might be high. Simulating TOCs from such variable
sources is next to impossible because TOC is not solely
the result of man-made organic chemicals. For sus-
pended solids, shape, porosity, surface-to-volume ratio,
charge and organic content (all or any), will impact sorp-
tion characteristics. None of these qualities are mea-
sured by the standard methods for measuring suspended
solids nor can they be reproduced in a simulated effluent.
In a simple system, such as reconstituted soft
water, it is reasonable to expect that for most chemicals a
doubling of the chemical concentration will double the
toxicity, at least in the effect concentration range. If the
solubility of the toxicant is being approached or there are
effects from water characteristics such as suspended
solids, then the toxicity might not double or conceivably
could more than double. For example, if a chemical with a
large n-octanol/water partition coefficient (log P) is largely
sorbed on solids, doubling the total concentration might
more than double the toxicity because the added chemi-
cal might remain in solution. Another important issue is
that equilibrium might not be established during the entire
test period and is probably unlikely to occur before the
test organisms are added. For example, in our TIE re-
search, we found various surfactants sorb to solids and
can be removed by filtration (Ankley et al., 1998). In these
experiments, however, filtration failed to remove surfac-
tants immediately after they were spiked in an effluent but
surfactants were removed after a few days equilibrium
time. Other chemicals are likely to show similar behavior
in regard to equilibrium time.
If several toxicants are involved, then their inter-
action (additivity, independent action, synergism) must be
measured or otherwise included in the confirmation pro-
cess (cf., Section 2). Since ratios might be as important
as concentration, the best way to spike when multiple
toxicants are involved is to increase each toxicant by the
same number of TUs (e.g., by doubling each). In this way
the ratios of the toxicities remain constant.
The fact that two or more toxicants fail to show
additivity is useful evidence in confirmation. Interpreting
spiking data might require a very high level of compe-
tence in both toxicology and chemistry; otherwise the
data could be very misleading. Using more than one
species of differing sensitivity is effective in adding confi-
dence to the results. When matrix effects are compli-
cated, other types of spiking can be done to reduce the
effects of the effluent matrix characteristics. If a method
exists for removing the toxicants from the effluent, such
as the C SPE procedures (EPA, 1993A), the extracts or
methanol fractions can be spiked with pure chemicals in
addition to spiking effluent, using the same principles as
described for effluents. The advantage in this approach is
that matrix characteristics such as suspended solids and
TOC will be absent or much reduced and will not affect
spiking experiments as much. The disadvantage is that
proof that the extracts or fractions contain the true toxi-
cants must be generated. Some approaches for doing
5-1
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this are given in Section 6. The use of the spiking ap-
proach is especially applicable to fractions from the C18
SPE column or the high performance liquid chromatogra-
phy (HPLC) column used for the isolation of non-polar
organics. In these procedures, the constituents are sepa-
rated from much of the TOC, suspended solids and
hardness, so that spiked additions might be strictly addi-
tive where they might not be in the effluent. Suggestions
and precautions about ratios and all other previously
discussed concerns apply here too. In addition, concerns
about the methanol percentages in the toxicity tests, the
amount of SPE or HPLC eluate required for the toxicity
tests and the issue of toxicity enhancement by methanol
must be considered in order to generate the appropriate
toxicity data. Spiking the methanol fractions with suspect
toxicants, however, does not provide the same confi-
dence about the cause of toxicity in the effluent as spiking
the effluent directly. The mass balance approach de-
scribed in Section 6 could be coupled with spiking the
effluent with a portion of the fractions to make the data
more relevant to whole effluent toxicity.
For chronic testing spiking a portion of the metha-
nol fractions, such as C18 SPE methanol fractions into
dilution water to mimic the effluent, requires some special
considerations as discussed in the chronic Phase I (EPA,
1992) and the new Phase II (EPA, 1993A). For any test
species, the effects of the methanol at the effluent spiking
concentration for the test species must either be essen-
tially non-existent or clearly established so that proper
interpretation is applied. The use of spiking for chronic
toxicants of the methanol fractions is not as easy as the
spiking for acute toxicants due to the limitations in the
quantity of methanol that would be added with each
fraction for the toxicity test. If the chronic toxicity effect
level is around or <25% effluent and the highest fraction
tested is 4x higher than the chronic effect level, add-back
tests can be conducted similar to the acute add-backs but
the quantity of methanol required for the testing and
analysis must be considered (cf., Section 2; EPA, 1993A).
As discussed in Phase II, once a suspect toxicant has
been tentatively identified, the steps of confirmation should
be started although sample volumes of methanol eluates
might limit the amount of testing (see Phase II, Section 2;
EPA, 1993A) with chronically toxic samples. Spiking of
appropriate levels for chronic toxicity for single chemicals
(or mixtures) is limited as sublethal data are not as
plentiful as acute data. The acute toxicity of some chemi-
cals might be altered by methanol (i.e., surfactants). The
possibility that this is occurring must be checked and a
correction applied if warranted. Spiking fractions also has
applicability for hidden toxicants; refer to Section 9 for
further details.
Spiking can also be done effectively when the
suspect toxicant(s) of concern can be removed. However,
since other toxicants might also be removed, the data
must be carefully interpreted. Ammonia is a good ex-
ample (cf., Phase II; EPA, 1993A) to use with this tech-
nique where one toxicant can be removed. Ammonia can
be removed from the effluent by passing samples over
the zeolite resin, after which the concentration can be
restored in the post-zeolite effluent by the addition of
ammonia. If toxicity is also restored, then it is likely that
there is sufficient ammonia to cause the toxicity observed.
However, it cannot be concluded from these data atone,
that ammonia is the cause of toxicity because the zeolite
can also remove substances other than ammonia. An-
other substance which is non-additive with ammonia yet
present at a lesser or the same number of TUs could
cause the initial effluent toxicity but not be discernable by
this removal technique. This is an example of a hidden
toxicant (see Section 9). For acute toxicity, zinc could
behave exactly this way because it is non-additive with
ammonia yet zinc is also removed by zeolite. Using other
ammonia removal methods, such as high pH stripping,
followed by spiking to the initial ammonia concentration
will enhance confidence that a hidden toxicant is not
present. Other examples involving the C18 SPE column
and various ion exchange resins would be approached
and interpreted similarly.
5-2
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Section 6
Mass Balance Approach
This approach is applicable only to those situa-
tions in which the toxicant(s) can be removed from the
effluent and recovered in subsequent manipulation steps.
The objective is to account for all toxicity to assure that
small amounts of toxicity are not being lost. This concern
is partly covered by the correlation approach (Section 2);
however, a totally different toxicant present at a small
concentration could appear as experimental variability in
the correlation and go unnoticed.
The mass balance concept is best described by
illustration for acutely toxic effluents and the C18 SPE
fractions. As described in Phase II (Section 2.2.7; EPA,
1993A) for acutely toxic effluents, the effluent has been
passed over a C SPE column which is then eluted with
the methanol/wafer fractions. After the toxicity tests on
the individual fractions are completed, add-back tests
can be initiated to determine whether all of the toxicity in
the original sample was accounted for in the SPE frac-
tions. For this step, there are three separate tests (with
dilutions and replicates to calculate effect endpoints) that
must be conducted which consist of the all-fraction test,
the toxic-fraction test, and the nontoxic-fraction test. As-
suming a complete recovery of all non-polar organics
from the SPE column, this should yield a solution of non-
polar organic compounds equal to the original sample
concentrations, In the mass balance approach, these
add-back tests are conducted using an aliquot of the
effluent that has passed through the C18 SPE column
(post-SPE column nontoxic effluent) or an aliquot of
dilution water. Each toxic fraction is added back to the
post-SPE column effluent, so that each is present at
original effluent concentrations (i.e., Ix effluent concen-
tration). For example for acutely toxic effluents, the toxic-
fraction test solution is prepared using methanol
concentrations as described in Phase II (i.e., Section
2.2.7; EPA, 1993A) and for each fraction where toxicity
was observed in the fraction toxicity test, 30 nl of each is
added to the same 10 ml of nontoxic post-C,, SPE
column effluent (or dilution water). A portion of each of
the remaining fractions where toxicity was not demon-
strated are now added to a second post-SPE column
aliquot at effluent concentrations for the nontoxic-fraction
test. Finally portions of all the fractions (e.g., n= 8 for
acutely toxic effluents) are added to a third post-SPE
columm aliquot at effluent concentrations for the all-frac-
tion test. If all the toxicity is exhibited in the toxic-fraction
test, then the all-fraction test results and the toxic-fraction
test results should be the same as in the unaltered
effluent. Results from the nontoxic-fraction test should
indicate that no toxicity is present. This mass balance (or
add-back) approach allows the researcher to ascertain
whether or not the toxicity in the toxic-fraction test equals
the effluent toxicity. Small amounts of toxicity can be
undetectable in the toxic-fractions when tested separately
or the toxicant(s) might not have been eluted from the C18
SPE columns. Unless mass balance experiments are
conducted, such loss of toxicity might not be detected. In
the effluent example discussed in Section 2, the toxicity
was contained usually in the 75%, 80%, and 85% frac-
tions and occasionally in the 70% fraction! The revalue,
slope, and intercept were all close to the expected values
if two toxicants (diazinon and CVP) were causing the
effluent toxicity (Figure 2-3). However, in Table 6-I the
results of mass balance tests indicate that toxicity from
the all-fraction test was greater than the toxicity of the
toxic-fraction test. While this difference is small, it did
seem to be real and was attributed to a small amount of
another toxicant in the 70% fraction. In 11 of 12 samples,
the results from the all-fraction tests indicate there was
greater toxicity than was found in the toxic-fraction tests.
On the few occasions when the 70% fraction was toxic, it
did not contain any of the three suspect toxicants. Without
the mass balance data, consistent presence of the addi-
tional toxicant would not have been discovered.
At the stage where the toxic-fractions have been
identified, the test of the fractions in a mass-balance test
is highly desirable. For chronic toxicity testing, the amount
of eluate available might be limited following the fraction
toxicity tests. Using eluate for the add-back tests might be
a trade-off between tracking toxicity and having sufficient
eluate to concentrate for further analysis. This limits the
add-back tests broad applicability for chronic toxicity Tl Es
unless the effluent is toxic enough that at 4x the chronic
effect level, the methanol concentrations do not exceed
^During development of the non-polar organic procedures, various
elution profiles were used that included the 70% methanol/water
fraction.
6-I
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Table 6-1. Comparison of Eff luent Toxicity and Toxicity Measured in
Effluent Fraction Add-back Tests
Sample
12/03/87
01/12/88
01/13/88
02/03/88-!'
02/03/88- 1 1
03/03/88-I"
03/03/88-M
03/23/88-I
03/23/88-H
04/28/88
05/17/88
05/17/88
Effluent
1.18
2.00
1.93
cl .00
2.00
1.15
1.33
3.70
2.86
2.27
2.27
2.27
Toxic Units (TUs)
All-fractions
1.64
2.94
2.86
1.15
1.75
1.06
1.52
3.03
2.86
1.72
2.04
1.67
Toxic-fractions
1.43
3.13
2.53
<1 .00
1.64
<1.00
1.13
2.86
2.44
1.64
2.00
1.59
Mean
2.13
2.18
2.00
"Values excluded from mean calculations due to less than values.
the organisms tolerance. For chronically toxic samples,
the all-fraction add-back test with C. dubia is not possible
due to high methanol concentrations in test cups unless
chronic toxicity is below 25% and add-backs are done
using 25% effluent as the high test concentration (cf.,
Phase II; EPA, 1993A). The data from the individual
methanol/water tests may be summed; however this ap-
proach must be considered more tentative than add-back
tests (see below).
A deficiency in the above approach to mass
balance is that there can be some toxicity in the post-SPE
column effluent which has not been removed by the C18
SPE but which is not present in concentrations high
enough to detect. The above mass balance approach
alone will not identify this. However, if the add-back tests
described above are repeated using a standard dilution
water, residual toxicity in the post-SPE column effluent
should cause the toxic-fraction test and all-fraction test to
show more toxicity when added to the post-SPE column
effluent than when added to dilution water. A confounding
effect of this approach is that if the toxicity is changed by
matrix effects (suspended solids or TOC), then the toxic-
ity will be different in the clean water test. Matrix effects
can be discerned, in part, by a third spiking experiment
where a portion of all of the fractions and a portion of each
toxic-fraction test are spiked into whole filtered effluent
(which has not passed through the C18 SPE column). If
the addback tests in dilution water indicates greater toxic-
ity than the addback tests with the post-SPE column
effluent, and the same type of addback test experiment
with filtered effluent (i.e., lum filter) indicate that the
fractions are exactly additive, then matrix effects are
indicated.
Some post-SPE column effluent samples develop
fungal or bacterial growth or perhaps a precipitate forms
after the effluent passes through the column. For the
fungal type of growth, this is thought to occur when some
methanol bleeds into the effluent as it passes through the
column and more rinsing will not eliminate this problem.
Some effluents consistently develop this type of growth in
the post-column effluent while others exhibit this pattern
in only an occasional sample. To alleviate this problem,
conditioning the column with acetonitrile has helped (cf.,
the acute Phase I (EPA, 1991 A) and chronic Phase I
(EPA, 1992) for details). When methanol fractions are
spiked into the effluent this problem might or might not be
enhanced; we have found this to be an effluent-specific
occurrence.
Caution is warranted in situations where toxicity
is contained in more than one SPE fraction. The re-
searcher should not necessarily expect the toxicity ex-
pressed by each individual fraction that is tested separately
to add up to the total effluent toxicity. First, toxicants may
not be additive and second, some toxicity which cannot
be detected in individual fractions may add to the whole
toxicity. For example, any one C18 SPE fraction may not
show toxicity but may contain some of the toxicant th§t is
in the adjacent toxic-fraction. In this case, the toxicity of
the toxic-fraction test would be less than expected. If this
happens in more than one pair of fractions, the sum of the
toxicity from the toxic-fraction test will be less than the
effluent toxicity or all-fraction test. These concerns are
especially important when several toxicants are involved
and one or more occur in more than one fraction.
For effluents where triallSPE column is not
used, but where the toxicants can fae removed from the
sample, the same objectives should be achievable, but
the methods will be different. For example, if an effluent
appears to contain a volatile toxicant, the mass balance
could be done on the trap and on the purged sample.
Since we have not yet done mass balance on samples
such as these we have no experience from which to offer
additional guidance or advice.
Some of the mass balance process begins in
Phase II, and there is a subtle difference in the purpose of
mass balances in Phases II and III. In Phase II, usually
only a few samples are used and mass balances are
necessary to determine the need for more identification in
those few samples. The mass balance is useful in early
stages of Phase II as well before toxicants are identified
at all, because it allows the investigator to decide if the
toxicants present at 2x or 4x whole effluent concentra-
tions are also expressing toxicity at lower concentrations.
In Phase III as many samples are tested, the
mass balance approach can provide information over
time with many samples whether or not the suspect
toxicants consistently account for all or the majority of the
toxicity. As illustrated above, the power of the mass
6-2
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balance approach to detect small degrees of toxicity is from the sample does not remove biologically non-avail-
better than for the correlation approach. able portions. An example of this situation may be the
alternative solvent extraction procedures which may re-
When a portion of the toxicant is not biologically move a bound toxicant(s) sorbed on suspended solids
available and therefore does not contribute to toxicity, with the solvent and is now toxic, yet it was not toxic in the
care must be taken to assure that removal of the toxicant unaltered sample.
6-3
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Section 7
Deletion Approach
In some situations, particularly for industrial dis-
charges, keeping the suspect toxicants out of the waste
stream influent or effluent for short periods of time and
also conducting toxicity tests on the wastewater simulta-
neously may be practical. When this approach can be
used, it offers the most convincing evidence obtainable
that the suspect toxicants are the true ones. Care must
be taken however, that other substances are not deleted
or that some characteristic such as pH does not change
also. If a researcher can be certain that all changes are
known, then this approach is definitive. Changes in the
toxicants with time are as much of a concern here as in
any other approach. These can be handled by the ap-
proaches outlined in earlier sections and the deletion
approach need not be done repeatedly; however, if it
were practical to do so, it would certainly be effective. If
some samples do not contain one or more suspect toxi-
cants, these effluent samples can be used to the advan-
tage in confirmation in much the same way as intentional
deletions described in this section can be used to confirm
toxicity.
7-1
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Section 8
Additional Approaches
This section mentions only a few of many steps
that can be used to further confirm the cause of toxicity.
The steps mentioned are mostly those that we have used
and found helpful and practical.
The pH is one of the most important effluent
characteristics that changes toxicity. The pH of POTW
effluents, sediment pore or elutriate waters, and ambient
waters will almost always rise when they are exposed to
air, especially in the small test volumes used in TIE work.
Commonly, pH in an effluent sample at 25°C will .rise
from 7.1-7.3 to 8.3-8.5 during a 24 h period. That pH
change is enough to increase ammonia toxicity (based
on total ammonia) about three fold. Such pH changes
can destroy work for some purposes, but by regulating
these pH changes, the pH fluctuations can be used to
great advantage for other purposes.
Phase II (EPA, 1993A) describes the use of pH
change to identify ammonia toxicity. The toxicity of some
metals, hydrogen cyanide and hydrogen sulfide among
others, is altered by pH change. Other characteristics,
such as hardness, can also be varied to see if the
changes in toxicity follow a predictable pattern. The
toxicity of some metals could be approached in this way.
Not all equilibria are as rapid as the ammonia equilibrium,
so the amount of time for equilibria to occur should be
controlled and standardized (cf., Phase II; EPA, 1993A).
Various time . periods may have to elapse before the
expected changes occur and this may differ with each
effluent. With the improved methods of pH control de-
scribed in the Phase I documents (EPA, 1991; EPA,
1992), much more use can be made of pH manipulation.
Often chemicals in effluent samples may not be
biologically available, and if they are not, then they are
not likely to cause toxicity. They may be made biologi-
cally available through some manipulation in Phase I and
subsequently identified in Phase II. Through confirma-
tion, the toxicity due to such a toxicant will become
apparent when the correlation indicates a poor fit (cf.,
Section 2). For many toxicants, biological availability can
be demonstrated by measuring body uptake. If the con-
stituent of concern enters the body from the effluent, it is
certainly biologically available. Exposure to pure com-
pounds may be necessary to establish which particular
organ should be evaluated for the toxicant. In acute metal
exposures using fish, most metals concentrate first in the
gills while non-polar organics concentrate in fatty tissues
such as the liver. When a chemical is metabolized by the
organism, a residue measurement for that compound is
not a valid measure of the lethal body burden because it
is unknown whether the metabolite is more of less toxic
than the parent compound. If the suspect toxicant has a
known mode of action, such as the acetylcholinesterase
inhibition produced by organophosphate pesticides, this
exposure effect can be measured to assess if toxic effects
conform with the predicted effect. The use of enzyme
blockers such as piperonyl butoxide (PBO) is also an aid
in confirming toxicity caused by specific classes of toxi-
cants (cf., Phase II; EPA, 1993A).
As additional steps are needed for confirming the
cause of toxicity, combinations of various Phase I and
Phase II procedures should always be used whenever
practical. When several results are combined and all
results are indicating the same type of toxicant, the data
are more conclusive than when only one procedure yields
predicted results.
Total dissolved solids (TDS) are a common prob-
lem in certain areas of the country and for certain indus-
tries. TDS will not cause toxicity from osmotic stress (this
can easily be shown because their toxicity is not related to
osmotic pressure) but rather TDS acts as a set of specific
toxicants. For toxicity caused by TDS, the ratios and
concentrations of the major cations and anions can be
measured analytically. A similar mix of these major ions
can be added to a dilution water to see if the expected
toxicity is present. By testing various mixtures, the re-
searcher can ascertain which of the TDS components
contribute most to the toxicity.
8-7
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Section 9
Hidden Toxicants
In the previous section, references were made to
the problem of hidden toxicants. Essentially there are two
situations which may produce the problem of hidden
toxicants. The first situation occurs when disparate ratios
ofTUs of two toxicants are present in the effluent sample.
Since the effect concentration is measured by diluting the
effluent, when disparate ratios occur, the TUs of the
toxicant present in fewer TUs in 100% effluent are so low
at the effect diluent, that its contribution if any, is not
measurable. This problem exists whether the toxicants
are additive or non-additive. This situation generally will
not be encountered in effluents that have very slight
toxicity (i.e., effect concentration 75% to 100%) because
little or no dilution is required to achieve the effect con-
centration. For those toxicants present in disparate ratios
in effluents with marginal toxicity, the chemical present at
the low levels may be nontoxic even in 100% effluent.
The second situation where hidden toxicant(s)
occurs is when the toxicants are non-additive or partially
additive in the effluent sample. These toxicants may
occur at approximately equal TUs or at disparate ratios of
TUs, as long as those present at lesser TUs are present
at 1 TU in the 100% effluent (cf., discussion of performing
correlation on these types of toxicants, contained in
Section 2).
If confirmation is being conducted for both acute
and chronic toxicity or if acute toxicity is being used as a
surrogate for chronic toxicity, the acute to chronic ratio
must also be considered. For example, consider an
effluent with toxicants A and B for which the acute-to-
chronic ratios are 3 and 12, respectively and the TUs for
acute toxicity are 2 and 1 in an effluent sample for A and
B, respectively. By definition, 1 acute TU (TU,) for toxi-
cant A equals 3 chronic TUs (TUo) and for B, 1 TU, = 12
TU,. In this example, the acute toxicity of the effluent will
be determined by A and the chronic toxicity will be
determined by B. If in another situation, the acute-to-
chronic ratios for two compounds were similar, then one
of the toxicants would determine the effect concentration
for both acute and chronic toxicity. These examples
illustrate the importance of acute-to-chronic ratios for
non-additive toxicants. Acute-to-chronic ratios have spe-
cial importance for additive toxicants when acute toxicity
is being used as a surrogate measure for chronic toxicity.
If acute toxicity is being used as a surrogate it must be
demonstrated that the cause of the acute toxicity is the
same as the chronic toxicity. When acute toxicity is used
as a surrogate for chronic toxicity in Phases I and II,
interpretation of the results can easily be biased and
these considerations are important.
When a toxicant can be removed from the efflu-
ent and recovered, the identification of the presence of a
hidden toxicant is more readily known. For example, the
use of the C SPE column may remove hidden toxicants.
The toxicantjs) is recovered in the eluate a&measured
both analytically and toxicologically. This type of hidden
toxicant may be observed if ammonia is present at con-
centrations that could cause toxicity. For example, in an
effluent sample ammonia is present at 3 TUs. Ammonia
will not be removed by the C SPE column and yet an
additional 1.5 TU of a non-pofar organic toxicant is evi-
dent when the C SPE eluate test is conducted. If the
discharger applied remedial treatment they would be able
to remove the ammonia toxicity yet the effluent would still
be toxic. The same concept of hidden toxicants can be
found when toxicants are removed by sublation which is
followed by recovery and concentration of toxicity (cf.,
Phase I; EPA, 1991 A; EPA, 1992). For example, sublation
can separate some surfactants, resin or fatty acids, and
polymers from such constituents as metals and ammonia.
Hydrogen sulfide can be removed by a purge and trap
method, thereby separating it from other effluent constitu-
ents.
Specific blockers of toxicity such as EDTA for
metals and PBO for organophosphates are also useful in
establishing the cause of toxicity. The more specific the
blocker, the more definitive are the results, However,
present knowledge does not allow us to be certain that
compounds such as EDTA do not also affect the toxicity
of other chemicals. Use of two specific blockers such as
EDTA and sodium thiosulfate for copper, allows more
definitive conclusions (cf., Phase I; EPA, 1992).
Manipulating characteristics such as pH is useful
but can easily mislead thinking. For example, if the efflu-
ent has ammonia toxicity, the toxicity due to ammonia
should disappear if the pH is lowered appropriately. These
results do not allow a conclusion that there are no hidden
9-I
-------
toxicants. If, however, the pH is lowered so as to eliminate
ammonia toxicity but the effluent toxicity exists or even
increases, then the likelihood of a hidden toxicant is high.
Unfortunately a complication to this rationale is that the
toxicity expressed at the lower pH may be totally artifac-
tual due to mechanisms of pH adjustments.
The best approach to find hidden toxicants is to
first use, those methods that alter the effluent the least,
can remove and recover removed hidden toxicants, and
are most specific for a few toxicants. This advice is most
applicable where the effort is to try to find out if some
specified type of toxicant is a hidden one, e.g., is there a
non-polar organic as a hidden toxicant.
If, however, the search is for any type of hidden
toxicant then every conceivable technique should be used
that would help to distinguish a hidden toxicant from the
suspect toxicant( Hidden toxicants are very hard to find
when ammonia is the primary toxicant. Various tests used
to identify ammonia as the toxicant, i.e., use of the zeolite
resin, graduated pH tests and air-stripping (EPA, 1993A),
all have a reasonable probability of changing the toxicity
of many other potential toxicants. For instance, it is known
that zeolite removes some non-polar organics and met-
als. Air-stripping (at pH 11) could also remove or destroy
many other chemicals as it often must be done for a
extended period of time to achieve good ammonia re-
moval. The graduated pH test results might also implicate
a metal as a toxicant (EPA, 1993A). If these tests were
conducted in Phase II (EPA, 1993A) and the results
consistently indicated ammonia toxicity, these .data indi-
cate that there are no hidden toxicants. The required
characteristics for a hidden toxicant to behave exactly as
ammonia are very specific and obtaining results like those
described above for a toxicant other than ammonia is
unlikely.
If the hidden toxicant is additive with the suspect
toxicant but occurs in a disparate ratio, the confirmation
effort must first emphasize confirming the cause of toxic-
ity (or remove the toxicity) of the primary toxicant. Then
toxicity from the hidden toxicant should be measurable.
The probability a hidden toxicant that has additive toxicity
will not express its toxicity using several Phase I or Phase
II techniques is less than the probability that a non-
additive toxicant will express its toxicity using several of
the same techniques.
If the remedial action for a primary toxicant is
specific and easy, such as a product substitution, the
search for hidden toxicants perhaps should be done after
the remedial action has reduced or eliminated the primary
toxicant from the effluent. The remedial action (especially
if it is treatment) may also eliminate the hidden toxicant(
What must be avoided if at all possible, is to carry out
expensive remedial action only to find that the effluent is
still toxic.
The problem of hidden toxicants is a major rea-
son a researcher should not accept the presence of toxic
concentrations of suspect toxicant as sufficient confirma-
tion (cf., Section 1). The presence of biologically unavail-
able forms (cf., Section 8) is a compelling reason not to do
so.
A thorough confirmation is resources well spent
in most instances. Non-additivity and disparate ratios
complicated by non-availability occur too frequently to by-
pass confirmation. Seasonal changes or changes without
a pattern, in effluent toxicants are further reasons to
perform the confirmation over a period of time to assure
that the entire suite of toxicants has been found.
9-2
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Section 10
Conclusions
Often the most laborious and difficult part of the
TIE is developing data to adequately establish the cause
of toxicity. In our experience, frequently the suspect
cause of toxicity is found without difficulty but developing
a convincing case to prove that the suspect cause is the
true toxicant is the challenge.
Especially for POTW plants, this confirmation
phase must be performed over a considerable period of
time to be certain that the cause of toxicity is not chang-
ing. TIEs on POTWs and some industrial categories are
not likely to be a one time event but will have to be
repeated as long as the inputs to the plant change. Our
current wastewater treatment plants were not designed to
remove specific chemicals, so there is no reason to
expect that they will remove everything which they re-
ceive. Especially where the control over the influent is not
complete, as is the case with POTW plants, a solid case
must be developed to assure that the cause of toxicity is
not changing.
10-1
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Section 11
When the Treatability Approach Has Been Used
As discussed in Phase I, two main approaches
may be used to remove a toxicity problem-toxicant iden-
tification and source control or treatability. Phases I and II
involve the first approach while treatability procedures
accompanied by toxicity testing are used in the second
(EPA, 19898; EPA 1989C).
In the second approach, treatment methods are
varied to determine which will remove toxicity without
identifying the specific toxicants. The treatability approach
requires as much confirmation as the toxicant identifica-
tion approach. Since the treatability approach should
remove toxicity, the confirmation procedures are some-
what different.
Repeat samples should be tested to ensure that
toxicity has been successfully removed. This should be
done over a sufficient length of time to assure that the
range of conditions are included during the confirmation
phase. Such events as seasonal changes, production
changes, storms, and intermittent operations all should
be included during the confirmation phase. Toxicity should
be consistently removed or appropriately reduced, as
required. Either acute or chronic toxicity removal can be
confirmed this way.
One must be absolutely sure that the toxicity to
resident species has been successfully removed. As has
been pointed out in Phases I and II, the effluent constitu-
ents producing toxicity to one species may not be the
same for other species. Toxicity by a given treatment
method may remove all toxicity for one species but not for
another. The species of concern must be tested in the
effluent from the treatment method selected. If chronic
toxicity is the concern, this testing may be more difficult
because chronic testing methods may not be available for
resident species. In selected cases, symptoms may be
substituted for the usual endpoints of chronic tests but
their use would be case specific.
11-1
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Section 12
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Drummond, R.A. and C.L. Russom. 1990. Behavioral
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